Transcript
Nutrient Elements Chapter 2 in Soils
Chapter 2
Nutrient Elements in Soils
Origin of Major Soil Constituents The nature of the soil parent material is usually the main influence on the amounts of the nutrient elements, other than N, present in a soil. The influence is due partly to the amounts originally present in the parent material and partly to the ability of the soil constituents to retain soluble forms of the elements against loss by leaching. With the large majority of soils, the parent material is inorganic and consists of fragments of weathered rock, together with the products of weathering, such as clay minerals, and oxides of Fe and aluminium (Al). However, with a small proportion of soils, the parent material is mainly organic, in the form of peat derived from past vegetation. The rocks from which most soil parent materials are formed may be igneous, metamorphic or sedimentary. Igneous rocks are of two main types, depending mainly on the temperature at which they solidified. Those that were formed at relatively low temperatures and pressures, such as granite, are generally acidic and are composed of a mixture of quartz, feldspars and muscovite micas (Table 2.1). On the other hand, igneous rocks that were formed at high temperatures and pressures are generally basic and contain a relatively high proportion of the ferromagnesian minerals, such as pyroxenes, amphiboles, olivines and biotite mica. The feldspars and muscovite micas have relatively high concentrations of Na and K, whereas the ferromagnesian minerals are relatively rich in Fe, Mg and Ca. The chemical structure of a mineral has a major influence on its behaviour during the processes of weathering. Thus quartz, which is composed entirely of linked silica tetrahedra, is stable and contains no elements other than silicon and oxygen. Feldspars, micas and the ferromagnesian minerals have more complex, less rigid structures and are less stable than quartz. Feldspars are basically Al silicates forming a three-dimensional framework and incorporating varying amounts of Na, K and Ca. Micas vary in composition but consist basically of a layer of Al–O octahedra, linked on each flat surface to a layer of Si–O tetrahedra, an arrangement often referred to as a 2 : 1 15
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Table 2.1.
Some primary and secondary minerals present in soils (from Paul and Huang, 1980).
Mineral group Primary minerals Silica Feldspars Micas Pyroxenes Amphiboles Olivines Phosphates Oxides
Name
Chemical formula of basic unit
Quartz Orthoclase Plagioclase Muscovite Biotite Augite Hornblende Forsterite Apatite Rutile
SiO2 KAlSi3O8 NaAlSi3O8 K(Si3Al)Al2O10(OH)2 K(Si3Al)(Mg,Fe)3O10(OH)2 Ca(Mg,Fe,Al)(Si,Al)2O6 (Ca,Na)2(Mg,Fe,Al)5(Si,Al)8O22(OH)2 Mg2SiO4 Ca10(F,OH,Cl)2(PO4)6 TiO2
Secondary minerals Kaolinite Clay minerals Chlorite Smectite/montmorillonite Vermiculite Oxides, hydroxides Goethite Haematite Gibbsite Allophane Pyrolusite Calcite Carbonates Dolomite Sulphur-containing Gypsum Pyrite
Si2Al2O5(OH)4 (AlSi3)AlMg5O10(OH)8 M*0.4(Al0.15Si3.85)(Al1.3Fe0.45Mg0.25)O10(OH)2.nH2O M*0.55(Al1.15Si2.85)(Al0.25Fe0.35Mg2.4)O10(OH)2.nH2O FeOOH Fe2O3 Al(OH)3 Al2O3.SiO2.nH2O MnO2 CaCO3 CaMg(CO3)2 CaSO4.2H20 FeS2
M*, variable metallic cation, such as K, Na, Ca.
lattice. Both feldspars and micas are negatively charged, and this is due largely to isomorphous replacement (i.e. the partial replacement of one element by another, during the formation of the mineral structure). For isomorphous replacement to occur, the difference between the elements in ionic radii must be less than about 15% and the difference in ionic charge no greater than 1. The most common type of isomorphous replacement is of Al3+ by Mg2+ and/or Fe2+. However, in some minerals, especially the ferromagnesian minerals, Fe2+ is often partially replaced by other metallic ions, including Co2+, Zn2+ and Cu2+ (Davies, 1994). In addition to any negative charge resulting from isomorphous replacement, a further negative charge may result from the dissociation of H+ ions at the edges of the lattice structure, a dissociation that varies with the ambient pH. The combined negative charge from both sources is balanced by cations, which are adsorbed on to the surface of the lattice structure and, in general, are exchangeable with other cations.
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When rocks are exposed to conditions at the earth’s surface, they are subject to both physical and chemical weathering. Physical weathering is due mainly to the expansion and contraction induced by changes in temperature, and results in the gradual disintegration of the rocks into smaller fragments. Chemical weathering is due to the interaction of rock material with water, oxygen and carbon dioxide (CO2). It increases with increasing acidity and therefore with increasing concentration of CO2. The rate of chemical weathering also tends to increase with increasing temperature, and is greater where new mineral surfaces have been exposed by physical weathering. The resulting changes in composition of the rock material include the partial loss of soluble constituents and the formation of new secondary minerals, such as clays, as shown in Fig. 2.1. Hydrolysis is a major component of chemical weathering and is involved, for example, in the transformation of feldspars and micas to clay minerals. Potassium feldspar is converted largely to the clay minerals, kaolinite and halloysite, with the release of silicic acid and free K+ ions. The weathering of muscovite involves, in addition to hydrolysis, the replacement of interlayer K+ by other cations, such as Mg2+, Ca2+ and Al3+, resulting in illite. Similarly, the weathering of biotite yields chlorite or vermiculite. As chemical weathering becomes more prolonged or intense, illite and vermiculite are transformed further to produce smectite, kaolinite and halloysite (see Fig. 2.1). In general, the clay minerals are composed of layers of Si–O tetrahedra and Al–O octahedra, linked in the ratio of either 1 : 1 or 2 : 1. Typically, there is some isomorphous replacement of Si4+ by Al3+ and of Al3+ by Mg2+ or SECONDARY MINERALS
PRIMARY MINERALS Olivines Pyroxenes Amphiboles
Non-crystalline hydrous oxides of Si, Fe, Al
Biotite
Chlorite
Goethite Haematite Gibbsite
Vermiculite
Muscovite
Illite
Feldspars
Non-crystalline aluminosilicates, e.g. allophane
Fig. 2.1.
Smectite (including montmorillonite)
Kaolinite Halloysite
Gibbsite
Weathering pathways of some common primary minerals (based on McBride, 1994).
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Fe2+, especially in the 2 : 1 lattice structures. Moderately weathered soils are often dominated by clay minerals that have a 2 : 1 lattice structure (e.g. smectite and illite), whereas more strongly weathered soils are dominated by clays with a 1 : 1 structure (e.g. kaolinite). Illite has the characteristic that it contains K+ ions between successive 2 : 1 layers, in positions in which the K is ‘fixed’ and therefore not readily exchangeable. Some clay minerals also incorporate small amounts of other metallic elements, such as Mn2+ and Zn2+, through the isomorphous replacement of Al3+ (West, 1981). Clay minerals, like micas, carry negative charges, due partly to isomorphous replacement and partly to the dissociation of H+ from the edges of the lattice structure, and these are balanced by adsorbed cations. The number of charges per unit weight of clay mineral, i.e. its cation exchange capacity, is an important property in relation to soil fertility and is often expressed in terms of milliequivalents (mEq) 100 g−1. In addition to clay minerals, the products of chemical weathering may include the insoluble oxides or hydroxides of Fe, Al and Mn, the carbonates of Ca and Mg, and various insoluble phosphates and sulphides (Davies, 1994). Although there is little N in primary minerals, some 2 : 1 clay minerals contain small amounts of N as ‘fixed’ ammonium within the crystal lattice. The quantity of any element that remains after weathering is influenced by a wide range of factors, including its abundance in the original rock, whether it behaves as a cation or anion, its ionic potential (valency/radius), and conditions such as pH and redox potential. During chemical weathering, there is normally some loss by leaching, and different constituents of the rock are lost differentially. As a result, the concentrations of nutrient elements in weathered and transported materials may differ markedly from the concentrations in the original rock. The monovalent cations, Na+ and K+, are lost more readily than the divalent ions, Mg2+ and Ca2+ and, amongst the common anions, chloride is lost more readily than sulphate, and sulphate more readily than phosphate. Past weathering sequences have resulted in the formation of sedimentary rocks, some of them consisting largely of clay minerals and others of fragments of igneous rock plus a cementing material, such as gypsum, calcite, silica and/or iron oxide. Shales are the most widespread type of sedimentary rock, while sandstones and limestones are dominant in some areas. The shales, although consisting mainly of clays, vary considerably in composition, especially in their proportions of clay- and sand-size particles and of OM. Limestones also vary, ranging from almost pure CaCO3 to dolomite limestone, which contains a mixture of MgCO3 and CaCO3, and with both types containing variable amounts of non-carbonate material. Sedimentary rock material has often been transported over long distances by water or ice, and this movement is partly responsible for the regional variation in soil parent material. Thus, much of the topography and the distribution of sediments in the northern hemisphere is due to the effects of the Pleistocene glaciation and the subsequent action of water. Wind erosion
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has also contributed to the distribution of sediments, resulting, for example, in the deposition of loess in the plains of central Europe and Asia. During the time span of geological activity, sedimentary rocks may have been subject to a number of cycles of weathering and deposition, or they may have been converted to metamorphic rocks by the extreme heat and pressure at depth below the earth’s surface. Slate, for example, is a metamorphic rock derived in this way from shale. As a result of differences in the extent of weathering and transport, the parent materials of soils range from slightly weathered igneous rock to sediments that have been weathered and transported repeatedly, and these differences result in wide variations in particle size distribution and in mineralogical composition. Within soils, the sand, silt and clay size fractions differ markedly in mineralogical composition. The sand and silt fractions (> 0.002 mm diameter) are both dominated by residual primary minerals, such as quartz, feldspars and micas, though the silt fraction may also contain a small proportion of secondary minerals. In contrast, the clay fraction (< 0.002 mm diameter) consists almost entirely of secondary minerals, such as clays and the hydrous oxides of Fe and Al (Paul and Huang, 1980). In temperate regions, hydrous Fe oxides are usually much more abundant than aluminium oxides, and they often occur as coatings on or interlayers between the crystalline clay minerals. Other elements are often associated with these hydrous oxides, due to the coprecipitation or adsorption of cations, such as Mn2+, Zn2+, Cu2+ and Co2+. There may also be some coprecipitation or adsorption of anions, such as phosphate, particularly under acid conditions. Chemical weathering continues after the formation of soil and often at an increasing rate, as microorganisms and plants release organic compounds, including acids, into the soil. Organic acids that form chelates with divalent and trivalent cations are particularly effective in releasing these ions from insoluble inorganic minerals. Although the nutrient elements released by weathering in soils are subject to uptake by plants, they are also susceptible to leaching, and this process tends, over many years, to deplete the surface soil of some nutrient elements.
Influence of Parent Material on Nutrient Elements in Soils The average concentrations of the nutrient elements in two igneous and three sedimentary rocks are shown in Table 2.2, and these differences are reflected to some extent in the composition of soils derived from the rocks. For example, soils derived from basalt tend to have higher concentrations of Mg and of the micronutrient cations than do soils derived from granite. However, although soils derive their initial supply of nutrient elements from their parent material, this influence tends to decline with time. The processes of leaching and gleying, and soil properties, such as pH and redox potential, all affect the extent to which the nutrient elements are retained in soils and the
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Table 2.2. Average concentrations (mg kg−1) of nutrient elements in two igneous and three sedimentary rocks (data from Mason and Moore, 1982; Krauskopf and Bird, 1995).
N P S K Na Ca Mg Fe Mn Zn Cu Co Cl I B Mo Se
Granite
Basalt
Shale
Sandstone
Limestone
13,159.007 13,390.007 13,158.007 45,100.007 24,600.007 9,900.007 2,400.007 13,700.007 13,195.007 13,145.007 13,113.007 13,122.4 13,170.007 < 0.03 13,121.7 13,116.5 13,120.007
13,152.03 13,610.03 13,123.03 5,300.03 16,000.03 78,300.03 39,900.03 77,600.03 1,280.03 13,186.03 13,110.03 13,147.03 13,200.03 < 0.03 13,115.03 13,120.6 13,120.3
1,1260.0 1,1700.0 2,400.0 26,600.0 9,600.0 22,100.0 15,000.0 47,200.0 13,850.0 13,195.0 13,145.0 13,119.0 13,180.0 13,122.2 13,100.0 13,122.6 13,120.6
– 05,170.05 05,240.05 10,700.05 3,300.05 39,100.05 7,000.05 9,800.05 10–100 05,116.05 1–10 05,110.3 05,110.05 05,111.7 05,135.05 05,110.2 05,110.05
– 015,400.28 1,200.08 2,700.08 015,400.08 302,300.08 47,000.08 3,800.08 1,100.08 051,120.08 051,114.08 051,110.1 051,150.08 051,111.2 051,120.08 051,110.4 051,110.08
forms in which they occur. Biological processes, such as the production of organic acids by plants and soil microorganisms, are also important, through their effects on soil pH and the ability of some organic acids to form soluble complexes or chelates with metal ions. Although the concentrations of many nutrient elements in soils are influenced by the nature of the inorganic parent material, N is an exception. Apart from the small amount of ammonium which may be fixed by clay minerals, almost all soil N is present in the OM and is derived ultimately from N2 in the atmosphere, by either biological or chemical fixation. Other elements, particularly P and S, are also present in soil OM, and their total concentrations in soils are therefore influenced by the amount of OM present. In general, the concentrations of most nutrient elements are higher in clay soils than in sandy soils, with loam and silty soils being intermediate: often the concentrations are inversely related to the proportion of quartz in the soil. Typical concentration ranges of 16 nutrient elements in the soils of temperate regions are shown in Table 2.3, together with the concentrations reported for three soils differing in parent material. Soils derived from peat are usually strongly leached and tend to have low concentrations of K, Ca, Mg and the micronutrient cations, but relatively high concentrations of N (McGrath and Loveland, 1992). Usually only a small proportion of the soil content of any nutrient element is available for uptake by plants at any one time, and this fraction is subject to continued depletion by uptake and leaching. However, it is
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Table 2.3. Typical concentration ranges of 17 nutrient elements in soils of temperate regions (based on Kilmer, 1979; Brady and Weil, 1999; plus data in subsequent chapters) mean concentrations of > 500 soils in USA (from Helmke, 1999) and concentrations in three agricultural soils from Scotland differing in parent material (from Ure et al., 1979).
N (%) P (%) S (%) K (%) Na (%) Ca (%) Mg (%) Fe (mg kg−1) Mn (mg kg−1) Zn (mg kg−1) Cu (mg kg−1) Co (mg kg−1) Cl (mg kg−1) I (mg kg−1) B (mg kg−1) Mo (mg kg−1) Se (mg kg−1)
Soils of temperate regions (range values)
Soils from USA (mean values)
Soil from granite
Soil from shale
Soil from sandstone
0.08–0.5 0.04–0.4 0.02–0.2 0.3–2.5 0.02–1.0 0.2–25.0 0.1–1.5 5,000–50,000 50–3,000 20–300 2–50 2–40 10–1,000 0.5–50. 2–100 0.2–5.0 0.1–2.0
– 18,000.026 18,000.120 18,001.50 18,000.59 18,000.92 18,000.44 18,000.026 18,330.026 18,048.026 18,017.026 18,006.7 – 18,000.75 18,026.026 18,000.59 18,000.26
– 18,000.09 18,000.06 18,003.90 18,001.60 18,000.36 18,000.14 10,000.02 18,450.02 18,036.02 18,007.0 18,001.2 18,034.02 < 0.1 18,003.6 18,001.8 18,000.09
– 35,000.12 35,000.10 35,001.50 35,001.20 35,000.13 35,000.96 35,000.02 35,920.02 347 35,006.0 35,019.0 35,160.02 35,002.5 35,022.0 35,000.7 35,000.12
– 35,000.04 35,000.02 35,001.70 35,000.71 35,000.53 35,000.39 20,000.24 35,440.24 35,057.24 35,008.5 35,004.0 35,500.24 35,000.9 35,009.9 35,001.1 35,000.24
constantly but slowly renewed by chemical weathering and, in some instances, by the mineralization of nutrients from OM and inputs from the atmosphere.
Inputs from the Atmosphere Several processes result in nutrient elements being transferred from the earth’s surface to the atmosphere and, as a consequence, being moved from one area to another before being deposited. Some of the transfers to the atmosphere occur naturally and some are the result of human activities. The natural processes include volcanic and geothermal activity, the formation of volatile compounds through chemical and biochemical reactions and the transport of dust by wind. Some elements are also transferred to the atmosphere from the sea surface, through the suspension and evaporation of droplets of sea spray. The processes that are due to human activity include the combustion of fossil fuels, the emission of pollutants from industrial processes and the increase in wind erosion caused by cultivation. Once in the atmosphere, gaseous molecules and solid particles are transported laterally for distances that may range from less than a millimetre to many thousands of
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kilometres. Eventually, almost all such molecules and particles are deposited on the land or sea surface, though possibly after chemical interaction with other molecules or particles in the atmosphere. Some of the deposition is ‘wet’, in rain or snow, and some is ‘dry’, occurring through the sorption of gaseous molecules by vegetation or soil or through the settling of aerosol particles by gravity. Wet deposition often supplies significant amounts of plant-available N, S and Cl, and sometimes Na and Mg, to soils. The amounts of Na, Cl and Mg deposited from the atmosphere are particularly influenced by transfers from the sea and therefore by proximity to the coast (Lag, 1987). In general, inputs by wet deposition are usually greatest in areas of high rainfall, while dry deposition is relatively more important in areas of low rainfall. The input of a nutrient by wet deposition reflects the amount of rainfall and the concentration of the nutrient in the rain, both of which can be measured fairly easily. However, the input by dry deposition is more difficult to assess, because both the concentration in the atmosphere and the average deposition velocity fluctuate with time and are influenced by the characteristics of a particular location. Thus, deposition velocity is influenced by particle size, by weather conditions and by the nature of the ground cover, including the type of vegetation (Sposito and Page, 1984). Assessments of the amounts of 16 nutrient elements deposited from the atmosphere at three rural locations in the UK are shown in Table 2.4. In industrial areas, much larger amounts of some elements, especially Zn and Cu (and several other non-nutrient elements, e.g. Cd), are deposited from the atmosphere and, if continued for many years, these may have a substantial Table 2.4. Deposition of nutrient elements at three rural sites in the UK: average annual amounts (kg or g ha−1 year−1) during the period 1972–1981 (Cawse, 1987).
N (nitrate only) (kg) S (sulphate only) (kg) K (kg) Na (kg) Ca (kg) Mg (kg) Cl (kg) Fe (g) Mn (g) Zn (g) Cu (g) Co (g) I (g) Mo (g) Se (g)
Range for three sites
Mean
36–44 43–58 9–20 21–41 15–19 5–11 42–74 1400–5700 90–200 480–1000 170–250 1.6–6.0 < 30–< 70 < 10 2.8–5.2
3,140.0 3,152.0 3,115.0 3,129.0 3,117.0 3,117.2 3,154.0 3100.0 3,135.0 3,660.0 3,220.0 3,113.3 < 50.0 < 10.0 3,114.0
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effect on the concentrations of the elements in soils (Sposito and Page, 1984). Even in the absence of marked industrial pollution, the total deposition of some elements, especially Fe, Zn, Cu and Co, may exceed the amounts removed by crops of silage or hay, as shown by comparing Table 2.4 with Table 3.5 (p. 55). Although, for most elements, much of the atmospheric deposition is not immediately plant-available, the addition to soil reserves may be important, especially where there is little release of nutrient elements by weathering.
Inputs from Fertilizers and Liming Materials In general, fertilizers are applied to grassland and other crops to meet the needs for N, P and/or K. However, it is unusual for the amount added in fertilizer to be taken up completely, and the remainder adds to the soil reserves, though there may be some loss by leaching and, with N, by volatilization. Also, a proportion of the fertilizer nutrient that is taken up is subsequently returned to the soil in dead plant residues and contributes to the soil reserves in this way. The repeated application of fertilizers, year after year, can therefore add appreciably to the total amounts of N, P and K in the soil, though the effect is generally greatest with P, which is least susceptible to loss. Nutrient elements other than N, P and K may also be supplied in fertilizers, as either intentional or incidental constituents. Sulphur is sometimes included deliberately in fertilizers and, in some regions, one or more of the micronutrient elements may be incorporated in order to overcome a potential deficiency. However, these elements and others, such as Ca, also occur as incidental constituents of fertilizers. For example, superphosphate contains approximately 11% S in the form of incidental calcium sulphate. Some fertilizers contain appreciable amounts of micronutrients, usually derived from the phosphate rock and/or potassium minerals used in their manufacture. The amounts vary widely, as illustrated in Table 2.5, largely because different sources of phosphate rock and potassium minerals have markedly different contents of the micronutrients. Liming materials, which are predominantly calcium carbonate or hydroxide, also contain small but variable amounts of nutrient elements, other than Ca, as incidental constituents (Table 2.6; Reith and Mitchell, 1964).
Recycling through the Decomposition of Plant Material The stems and roots of grassland plants have a maximum lifespan of a few years, and individual leaves and roots often survive for only a few weeks or months. Even when some of the herbage material is removed by cutting or grazing, a substantial proportion of the herbage, plus all the root material, is eventually returned to the soil through death and decomposition. The
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Table 2.5. Reported concentrations (mg kg−1) of some micronutrient elements in ammonium nitrate, superphosphate and potassium chloride fertilizers: (i) Verloo and Willaert, 1990; (ii) Adriano, 1986; (iii) Raven and Loeppert, 1997.
Fe Mn Zn Cu Co B Mo
Ammonium nitrate (i)
(i)
(ii)
(i)
(ii)
(iii)
192.7 39.7 11.7 3.6 3.3 – –
1150 9 244 23 15 – –
– 890 165 15 77 132 335
145.7 < 0.7 6.1 1.7 22.5 – –
– 3.5 10.7 3.1 22.7 16.7 26.7
440.7 2.8 4.6 2.7 < 0.07 – –
Superphosphate
Potassium chloride
Table 2.6. Concentrations of 13 nutrient elements in agricultural limestones from the USA (from Adriano, 1986).
P (%) S (%) K (%) Na (%) Ca (%) Mg (%) Fe (mg kg−1) Mn (mg kg−1) Zn (mg kg−1) Cu (mg kg−1) Co (mg kg−1) B (mg kg−1) Mo (mg kg−1)
Mean
Range for 194 samples
4,110.02 4,110.11 4,110.23 4,110.03 4,130.2 4,114.9 4,300.71 4,330.71 4,131.71 4,112.7 4,1< 1.71 4,114.71 4,111.1
0.001–0.56 < 0.01–1.35< < 0.001–1.80< < 0.001–0.15< 18.0–39.8 0.04–12.9 31,100–31,000 20–3,000 < 1–425< < 0.3–89< . 1–6 < 1–21< < 0.1–92 < .
nutrient elements in the dead herbage and root material are thus recycled and, if there is little input from external sources, this recycling provides the main source for new plant growth. The decomposition of plant residues depends mainly on microbial activity in the soil, and recycling is therefore more rapid in warm moist conditions and when the soil pH is in the range 5–7. Decomposition is also increased by the activities of soil fauna, such as earthworms and insects, which transfer some of the dead plant material to below the surface of the soil, thus making it more accessible to the soil microorganisms. Earthworms and insects also consume dead material, which, after excretion, is more susceptible to further decomposition by microorganisms. Increasing the productivity of a grassland sward – for example, by
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promoting the growth of clover or by the use of fertilizer N – tends to increase the population of earthworms and, in highly productive grassland, the weight of earthworms per hectare may equal the weight of grazing livestock (Haynes and Williams, 1993). During the decomposition of plant residues, those nutrient elements that are constituents of organic molecules are converted to inorganic forms. This process, known as mineralization, is due mainly to extracellular enzymes released by the soil microorganisms, and results in the release of N as ammonium, P as phosphate and S as sulphate ions. At the same time, because soil microorganisms have relatively high concentrations of these nutrient elements, a proportion of the amount mineralized is immobilized in the microbial biomass. The balance between mineralization and immobilization depends mainly on whether the size of the microbial biomass is increasing or decreasing: when it is increasing, there is inevitably some increase in immobilization but, when the microbial biomass is decreasing, there are increases in the amounts of N, P and S mineralized, due to release from the dead microbial tissue. The mineralization and immobilization of the nutrient elements in plant residues do not necessarily match the mineralization and immobilization of carbon, which may undergo net mineralization to CO2 at the same time that N, for example, is undergoing net immobilization. Such differences are due largely to differences in composition between the plant material and microbial biomass. Plant residues typically have a C : N ratio of > 20 : 1, whereas the microbial biomass has a ratio of about 10 : 1. Similarly, the C : P ratio of plant litter is often > 250 : 1, while that of microbial biomass is about 30 : 1 (Clark and Woodmansee, 1992). The amounts of herbage material that decompose in situ in grassland swards are difficult to measure, partly because some of the material is consumed during the senescent stage by soil fauna and partly because the rate of decomposition varies widely with factors such as temperature and rainfall. In a semi-natural chalk grassland in the UK, not fertilized and not grazed by farm livestock, grass leaves were estimated to live for periods varying from 6 weeks to 6 months, depending on the species and the time of year (Williamson, 1976). With white clover, little of the stolon material appears to survive for more than one growing season (Marriott, 1988). Physiological measurements carried out on intensively managed grass swards indicated that the amount of herbage decomposing in situ each year was broadly similar to the amount consumed by grazing animals or harvested for hay or silage (Parsons, 1988). With swards that are grazed lightly or are cut infrequently, the proportion of the herbage decomposing in situ is higher. Leaves senesce more rapidly during periods of extreme temperatures and of drought, but some leaf material dies throughout the year. Damage caused by grazing animals also increases the death of leaves and, in general, the proportion of the herbage that decomposes in situ is probably about 20% greater when a sward is managed by grazing rather than cutting (Hassink and Neeteson, 1991). Overall, depending on the intensity of management and other factors,
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the total amount of herbage material decomposing in situ varies from less than 1000 DM ha−1 year−1 in some natural and semi-natural grasslands to more than 10,000 kg DM ha−1 year−1 in heavily fertilized swards. In some natural grasslands, particularly those on acid and/or poorly drained soils, there is a tendency for dead herbage and stem material to accumulate as a partially decomposed mat at the soil surface. Such an accumulation inevitably reduces the rate at which the nutrient elements are recycled. Assessing the amount of root material that decomposes in grass swards is also difficult. A newly sown grass sward may produce about 3000 kg root OM ha−1 in 6 months, increasing to 10,000 kg ha−1, including both living and dead roots, after 3 or 4 years. However, in dry conditions, the amount of roots may be only 5000 kg or less after several years (Haas, 1958). Under long-term grassland, the total amount of root material is often between 10,000 and 20,000 kg ha−1 and, in prairie grassland, it may be as much as 25,000 kg ha−1 (Black and Wight, 1979). Although there is a lack of information on the length of life of grass roots, there is some evidence that, when swards are defoliated at intervals of 3–4 weeks, the average length of life is about 5–6 months (Garwood, 1967a; Troughton, 1981). Certainly, the death of roots is increased by defoliation (Evans, 1973; Eason and Newman, 1990). When defoliation is infrequent, roots may live for more than a year, and the average length of life of grass roots in prairie grassland has been estimated at between 2 and 5 years (Dahlman and Kucera, 1965; Clark, 1977; Redmann, 1992). Soil temperature may also have an effect, as illustrated by the finding, with young perennial ryegrass, that the length of life of the roots declined as temperature increased over the range 15–27°C (Forbes et al., 1997). In grass–clover swards, the clovers produce much less root material than do the grasses (Young, 1958), but clover roots decompose more quickly than grass roots (Laidlaw et al., 1996). From these various observations, it appears that the annual turnover of root material probably ranges from less than 1000 kg DM ha−1 in some natural and semi-natural grasslands to more than 8000 kg DM ha−1 in some intensively managed swards. Evidence from radiocarbon dating indicated that the annual turnover of OM, mainly roots, in an old grassland sward in lowland UK was 4000–5000 kg ha−1 (Jenkinson, 1969). Even where the grass species are mainly annual, as in parts of California, the amount of root material undergoing decomposition each year may be as much as 4500–7500 kg ha−1 (Jones and Woodmansee, 1979). In general, the amounts of both herbage and root material undergoing decomposition each year increase with increasing intensity of management, and this implies a corresponding increase in the amounts of nutrient elements being recycled. However, in most grassland soils there is a tendency for the content of OM to increase, and this trend continues for many years if the sward remains undisturbed. Consequently, with N, P and S, which are inherent constituents of soil OM, a proportion of the amount returned each year accumulates in the OM and is not released for
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renewed plant uptake. Only when the equilibrium content of soil OM is attained does the annual amount of N, P or S returned in dead plant residues (plus animal excreta if the sward is grazed) equal the amount that is potentially available for uptake. However, whether or not there is an accumulation of OM, the uptake of nutrients by plant roots and their return in dead herbage and roots are important in counteracting the tendency towards nutrient depletion caused by leaching. Where there is an accumulation of soil OM, the amounts of nutrient elements in circulation may increase slowly from year to year, due to small inputs from the weathering of soil parent material and/or from the atmosphere. On the other hand, if a grassland sward is ploughed and resown, the OM content of the soil decreases, due to an increased rate of decomposition, and, consequently, there is an increase in the release of nutrients in soluble forms. These soluble nutrients may be taken up by plants but, if not, are susceptible to loss by leaching.
Recycling through the Excreta of Grazing Animals The presence of grazing animals increases the rate at which nutrient elements are recycled, because the animals utilize for live-weight gain or milk production only a small proportion of each of the nutrients that they consume, and nutrients that are not utilized are returned to the soil in dung and urine. The nutrients in dung are mainly insoluble and become available slowly, whereas those in urine are water-soluble and are therefore potentially available for plant uptake immediately after deposition. The extent to which the recycling of nutrients in grassland is increased by the presence of grazing animals is, of course, influenced by the density of stocking, by the proportion of their time that the animals spend in the field and by the concentrations of nutrients in the herbage consumed. Dairy cattle are usually taken from the field twice per day for milking, and some of their excreta are deposited on farm tracks and in buildings, with the result that the recycling of nutrients is incomplete. Often, 10–35% of the excreta are deposited in milking sheds or on yards and roadways (Nguyen and Goh, 1994a). However, dairy cattle are often given supplementary feed in addition to grazed herbage, and there may then be a net addition of nutrients to the soil through the return of excreta. The proportions of the grazed area that are affected by dung and urine during any 1 year will depend on the density of stocking, the type of stock and the daily production of dung and urine. However, for dairy cattle grazing at an intensity of 700 cow-days ha−1 year−1, assuming an average of 14 faecal excretions per day with 85% in the field being grazed, each covering an area of 0.07 m2 (see Table 4.12, p. 93), then about 6% of the grazed area would be actually covered by dung each year. However, a larger area, perhaps 20–30% of the total, would be affected by its proximity. With urine, assuming a urination frequency of ten times per day and 85% of the
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urinations in the field being grazed, plus an average area per urination of 0.35 m2 (see Table 4.12) with no overlapping, then urine would be deposited on 21% of the area each year. However, the area of the sward actually covered by a single urination is influenced by a number of factors, including soil moisture content, the presence of macropores that open to the soil surface and the amount of herbage present (Williams and Haynes, 1994). Whatever the area covered by urine, a larger area is affected due to the lateral diffusion of nutrients and the lateral component of root growth, and this may extend to two to five times the area actually covered (see Table 4.12; Lantinga et al., 1987). The recycling of nutrients in excreta, and especially in urine, differs to some extent between cattle and sheep, even though the concentration of nutrients in urine is similar for sheep and cattle if they consume the same pasture (Williams and Haynes, 1993). Cattle produce larger volumes of urine but usually at less frequent intervals (see Table 4.12), and one effect of the larger volume is that the urine penetrates to a greater depth in the soil and tends to result in a greater concentration of nutrient ions, such as NH 4+ and K+, in the soil solution of the urine patch. With sheep, the urine patches are smaller and more evenly spread, resulting in a lower concentration of the nutrient ions in the soil solution. Studies with artificially applied urine showed that, typically, cattle urine moved to a depth of 400 mm by flow through macropores, whereas, with a sheep urination, the depth was only about 150 mm (Williams and Haynes, 1994). Sometimes urine scorches the grass on which it is deposited and the reduction in growth, though temporary, results in a reduced uptake of the nutrients contained in the urine. Scorching is most likely to occur in dry conditions and especially when no rain falls in the days following deposition, and when there is little if any accumulation of dead leaf, stubble and root material at the soil surface (Williams et al., 1999). The visible effect of excreta on grass growth is due mainly to the N and K in urine, and usually this is most apparent about 2–3 months after deposition, though the effect of K in particular may continue for longer (Wolton, 1979). There is often no visible effect of the other nutrient elements in either dung or urine, though there may be a significant impact in the long term. With moderately intensive management, the areas actually covered by urine typically receive the equivalent of about 500 kg N ha−1 and 500 kg K ha−1 (see Table 4.13, p. 93) and, in some situations, the rates of return are considerably higher. A rate equivalent to about 1000 kg K ha−1 was thought to be typical of productive pastures in New Zealand (Saunders, 1984). Since both the N and K are in readily available forms, the inputs to a urine patch exceed the amounts that would be taken up by grass during a period of several months, and they are therefore susceptible to loss. Depending on rainfall, substantial losses of both N and K can occur from grazed swards by leaching and, whatever the rainfall, appreciable amounts of N can be lost as gases through the volatilization of ammonia, denitrification and nitrification. Losses of nutrients through leaching are reduced when the soluble forms in the soil are distributed as evenly as possible over the area and, with a grazed
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sward, a more even distribution of nutrients is achieved by a high density of relatively small animals than by a lower density of large animals. The effect of urine on uptake by plants is often more long-lasting with K than with N, since the N is lost more readily through leaching and volatilization. In a sward that was intensively grazed by sheep, the proportion of the grazed area affected at any one time by K was estimated to be about three times greater than that affected by N (Morton and Baird, 1990). With moderately intensive management, the patches actually covered by dung might well receive the equivalent of about 1000 kg N, 350 kg P and 380 kg K ha−1 (see Table 4.13). Most of the N and P in dung is in relatively unavailable forms, and some of the N that is potentially available is lost through the volatilization of ammonia. The herbage actually covered by dung may be smothered temporarily, but growth is often stimulated in the area immediately surrounding the patch (MacDiarmid and Watkin, 1971; Weeda, 1977), and the area affected is often about three to five times the area covered (MacDiarmid and Watkin, 1972; During and Weeda, 1973) or up to 15 cm from the edge of the dung patch (Deenen and Middelkoop, 1992). The increase in growth is usually most apparent 2–3 months after the deposition, but the visible effect is accentuated because herbage in the affected area is rejected by grazing animals while visible dung is present (Marsh and Campling, 1970). Sometimes a positive effect on growth may continue for up to 2 years (Haynes and Williams, 1993). There is often a rapid increase in the amount of microbial biomass in dung during the first 5 days after deposition and, during this period, there is a substantial mineralization of organic N (Yokoyama et al., 1991). The decomposition of dung occurs most rapidly in moist conditions (Dickinson and Craig, 1990) and is accelerated by the presence of soil fauna, such as earthworms and dung beetles. Earthworms are particularly important in cool temperate regions, at least when their presence is not curtailed by a lack of moisture (Holter, 1979), whereas dung beetles are most prevalent in warm climatic conditions, such as those in Australia (Tyndale-Biscoe, 1994). The effect of soil fauna in burying and fragmenting dung is reflected in a greater reutilization of the nutrients by grassland plants when such fauna are present (Fincher et al., 1981). Rainfall after the deposition of dung also increases its rate of decomposition (MacDiarmid and Watkin, 1972), as do other causes of fragmentation, such as birds searching for insects. In general, the rates of fragmentation and decomposition are greater for sheep dung than for cattle dung. In a study in New Zealand, sheep dung deposited in spring remained visible on the sward surface for only a few weeks, whereas cattle dung remained for about 12 months (Williams and Haynes, 1993b). As dung and urine are distributed unevenly over grazed grassland, there is always a large spatial variation in the nutrient status of the underlying soil, variation which resembles a mosaic whose pattern is slowly changing. The change in the pattern is largely due to the decline in the amounts of nutrients under an individual dung or urine patch, resulting partly from plant uptake
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and partly from losses through leaching or volatilization. Sometimes the dung and urine patches are distributed fairly evenly within the grazed area (MacDiarmid and Watkin, 1972), but there is often spatial variation on a larger scale, due to the tendency of animals to excrete in certain areas, such as near water troughs or in shaded or sheltered areas (West et al., 1989; Wilkinson et al., 1989; Mathews et al., 1994). Sheep have a greater tendency than cattle to ‘camp’ and excrete in such areas (Wilkinson and Lowrey, 1973; Nguyen and Goh, 1994a). Whatever the pattern of distribution of the excreta, the non-uniform return often results in some parts of a grazed sward being deficient in one or more nutrients, while other parts have an excess (West et al., 1989). However, there is little information on the magnitude of this problem under different grazing systems or on the lateral and vertical distribution of excreted nutrients in the soil and the changes that occur with time. In general, the spatial variation in nutrient status tends to be minimized by grazing with small rather than large animals and by measures that discourage the animals from congregating repeatedly in certain areas. Spatial variation is also reduced when animals produce relatively dilute urine at relatively frequent intervals. With dung, the effects on spatial variation tend to be less when there is an active population of earthworms or dung beetles, and variation can also be reduced by harrowing. When livestock are housed, their excreta are normally collected and stored as slurries or manures before being applied to the land. Slurries contain both dung and urine, mixed with a variable amount of water, mainly from washing the floors of the housing. Farmyard manure consists largely of dung, with a variable amount of urine and bedding material, such as straw. Some average concentrations of nutrient elements in cattle manures and slurries are shown in Table 2.7. Losses of N can occur from both slurry and farmyard manure by volatilization and, partly for this reason, the ratios of N : P and N : K in slurries and manures are usually lower than those required by plants. Consequently, when slurries are applied to grassland at rates chosen to supply adequate N, the amounts of P and K may be excessive. Thus, in a study in which three rates of cattle slurry were applied annually to grassland for 16 years, with no application of inorganic fertilizers, there was a substantial accumulation of P and K in the surface soil, though little effect below 10 cm (Christie, 1987). In order to achieve the maximum utilization of nutrients when slurry is applied in conjunction with fertilizers, the rate and timing of each application should be adjusted to take into account the effects of the others (MAFF, 1994).
Inputs from the Disposal of Sewage Sludges Sewage sludge, which is a major waste product of urban communities, often presents a disposal problem, but it does contain appreciable amounts of nutrient elements, especially N and P. The amounts of K are relatively small,
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Table 2.7. Some reported average concentrations of nutrient elements in cattle manures and slurries (% or mg kg−1 in DM): (i) 400 samples, Maryland (Brady and Weil, 1999); (ii) monthly samples, 1 year, seven farms, North Carolina (Safley et al., 1984); (iii) samples from six farms (Levi-Minzi et al., 1986); (iv) average values (Van Dijk and Sturm, 1983); and (v) 13 farms, north-west Spain (Diaz-Fierros et al., 1987).
N (%) P (%) S (%) K (%) Na (%) Ca (%) Mg (%) Cl (%) Fe (mg kg−1) Mn (mg kg−1) Zn (mg kg−1) Cu (mg kg−1) B (mg kg−1)
(i) Manure, dairy cow, USA
(ii) Manure, dairy cow, USA
2.4 0.7 0.3 2.1 – 1.4 0.8 – 1800.4 165.4 165.4 30.4 20.4
4.7 0.95 – 2.18 0.41 1.17 0.60 1.00 1400.41 190.41 200.41 40.41 –
(iii) (iv) (v) Manure, beef Slurry, cattle, Slurry, dairy cattle, Italy The Netherlands cattle, Spain 2.70 0.55 – 3.37 – 1.4 0.6 – 1220.41 150.41 143.41 49.41 –
5.00 1.05 – 5.67 – 1.86 0.90 – – – – – –
4.00 0.70 – 4.15 0.98 1.71 0.70 – – – – – –
as most of the K remains in the liquid effluent. However, sludges vary in their composition, depending on factors such as the source of the sewage, the method of treatment and how they are stored before application to the soil (Oberle and Keeney, 1994). The simplest method of sludge treatment involves screening to remove fragments of metal and glass, followed by settlement of the solid particles in water to produce raw or primary sludge. Raw sludge is often treated by the activated sludge process, producing a secondary sludge which contains a large proportion of bacterial material. Although this secondary sludge can be applied to land, it is often treated further, in order to reduce the bulk of the sludge and to reduce odour and the population of pathogenic organisms. Anaerobic digestion at about 35°C for 5–35 days is widely used. Until recently, most sewage sludge was disposed of either at sea, in landfill sites or by incineration. However, it is now increasingly being disposed of in ways that are less damaging to the wider environment and, in particular, by application to agricultural land. This method of disposal has the advantage that nutrient elements are returned to the soil, but there are potential disadvantages if the sludge contains toxic materials or pathogens, or if it has to be transported over long distances. Grassland is often more suitable than arable land for the disposal of sludge, because grassland areas are more readily available throughout the year and because the presence of the grass reduces the compaction of the soil caused by vehicles applying the sludge. There is also less risk of surface runoff from grassland. In contrast,
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the use of arable land is restricted to those times of the year when crops are absent and when there is little risk of soil compaction (Smith, 1996). Some data on the concentrations of nutrient elements in sewage sludges are shown in Table 2.8. In comparison with typical concentrations in soils (Table 2.3), sewage sludges usually contain much higher concentrations of Cu and Zn, and appreciably higher concentrations of N, P, S, Ca, Mo and Se. Of the toxic materials that may be present in sewage sludge, the most important are the metallic elements that are derived from industrial wastes, particularly Zn, Cu, Cd, Cr and Ni. The actual amounts clearly depend on the proportions of industrial and domestic wastes in the sludge and on the types of industry in the locality. A high proportion of each of these metallic elements is bound initially to the OM of the sludge and, although the metals are released slowly in available forms as the sludge decomposes, they tend to be retained in the soil. Repeated applications may therefore result in an accumulation of metallic elements in the soil, which may ultimately have toxic effects on plants or on livestock consuming the plants. Some metallic elements may also have a harmful effect on microbial processes, such as nitrification, but, in general, the soil microbial population appears to adapt to concentrations in the soil that are less than those that are toxic to plants or livestock (Smith, 1991). However, since the metallic elements in sewage sludge tend to accumulate in the surface layers of the soil, there may
Table 2.8. Mean and range concentrations of nutrient elements in sewage sludges as reported in four investigations (% or mg kg−1 DM): (i) 16 samples, cities, USA (Furr et al., 1976); (ii) > 250 samples, north-east USA (Sommers, 1977); (iii) 42 samples, UK (Berrow and Webber, 1972; and, for I, Whitehead, 1979); and (iv) 555 samples, UK survey (Smith, 1996). (i) 2 ,132.9 (1.5–5.8) N (%) 2 ,131.6 (1.0–2.7) P (%) – S (%) 2 ,131.2 (0.3–3.9) K (%) 2 ,130.4 (0.08–1.5) Na (%) 2 ,133.6 (0.8–11.6) Ca (%) 2 ,130.6 (0.2–1.13) Mg (%) 2 ,130.4 (0.05–1.02) Cl (%) 2 ,133.1 (0.9–8.3) Fe (%) Mn (mg kg−1) 2,190 (30–520) Zn (mg kg−1) 2,130 (560–6,890) Cu (mg kg−1) 1,350 (580–2,890) Co (mg kg−1) 2,110 (3.7–17.6) I (mg kg−1) 2,137.8 (1.0–17.1) B (mg kg−1) 2,137 (16–90) Mo (mg kg−1) 2,114 (1.2–40) Se (mg kg−1) 2,133.1 (1.7–8.7)
(ii)
(iii)
(iv)
2,133.9 (0.1–17.6) 2,132.5 (0.1–14.3) 2,131.1 (0.6–1.5) 2,130.4 (0.02–2.6) 2,130.6 (0.01–3.1) 2,134.9 (0.1–25.0) 2,130.5 (0.03–2.0) – 2,131.3 (0.1–15.3) 2,380 (20–7,100) 2,790 (100–27,800) 1,210 (80–10,400) 2,135.3 (1–18) – 2,177 (4–760) 2,128 (5–39) –
– – – – – – – – 2,132.4 (0.6–6.2) 2,500 (150–2,500) 4,100 (700–49,000) 2,970 (200–8,000) 2,124 (2–260) 2,135.7 (0.9–17.4) 2,170 (15–1,000) 2,137 (2–30) –
– – – – – – – – 2 ,131.6 (0.2–10.7) 2,380 (55–1,390) 1,140 (280–27,600) 2 ,590 (70–6,140) 2 ,110 (< 2–600) – 1,130 (15–1,000) 2 ,135 (< 2–154) 2 ,133 (< 2–15)
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be a risk to grazing animals, not only through the consumption of herbage but also through their direct ingestion of soil. Many countries have introduced restrictions to prevent the excessive accumulation of metallic elements in the soil – for example, by specifying: (i) minimum periods between the application of sludge and the grazing or harvesting of grassland; (ii) maximum permissible concentrations in soils; and/or (iii) maximum permissible average rates of addition over a 10-year period (MAFF, 1993; Kabata-Pendais and Adriano, 1995). In the UK, it is permissible for livestock to be grazed on sludge-treated grassland after a minimum period of 3 weeks, but this restriction still entails some risk of the animals consuming sludge adhering to the herbage or present in ingested soil (Smith, 1996).
Availability of Nutrient Elements to Plants in Relation to Chemical Form and Soil Processes With any nutrient element, the fraction immediately available for plant uptake is that present in the soil solution. For most elements, this soluble fraction is largely in simple ionic form, but a variable proportion may occur in small molecular organic compounds. For example, some N may occur in amino acids, some P in inositol phosphates and a proportion of the divalent metallic ions in various organic complexes. Many nutrient elements are able to occur in the soil solution in more than one chemical form, i.e. they have a mixed speciation, and the speciation may vary with factors such as pH, redox conditions and the presence of organic complexing agents. However, the simple ions are generally taken up more readily than the organic compounds. The fraction of any nutrient present in the soil solution is usually only a small proportion of the total, but, when this fraction is depleted by plant uptake and leaching, replenishment tends to occur through a number of natural processes. These include the slow weathering of soil minerals, the mineralization of OM, inputs from the atmosphere, ion exchange and equilibria between soluble and insoluble inorganic compounds. With some elements, particularly N and S, the mineralization of OM is the process that has the greatest influence on the available supply. However, with other elements, such as K and Mg, the available supply is strongly influenced by the cation exchange capacity of the soil and hence by the amount and nature of the clay minerals. With a number of elements, particularly the metallic micronutrients, the formation of soluble complexes with organic molecules is important. Not all soil processes tend to increase the solubility of nutrient elements; some have the opposite effect. Thus, an element released in soluble form in the soil solution may be adsorbed or occluded in a non-exchangeable form by soil constituents such as hydrous Fe oxides; it may become insoluble by reacting with other soluble compounds, or it may be immobilized in organic material, particularly the microbial biomass. Nutrients in these insoluble forms are currently unavailable, though they may become more soluble
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and available again over a period of time. With the divalent and trivalent cations that form soluble organic complexes, the formation of a complex tends to increase the amount in solution, but it increases plant uptake only if the complex is absorbed intact, or if it is dissociated readily at the root surface. The nutrient elements in grassland soils, and especially their available forms, are usually present in highest concentration in the top few centimetres. This is largely because leaf litter and animal excreta are deposited and fertilizers are applied on the surface of grassland, and the nutrients are moved downwards only by water or by soil fauna. Within the soil, the accessibility of nutrient elements to plant roots is influenced by the extent to which the soil particles form aggregates, by the size and distribution of pores between the aggregates and by the relative amounts of soil solution and air in the pore space. Fertile soils generally have a large proportion of their aggregates with a diameter of between 1 and 5 mm, and roots grow and take up water and nutrients mainly from the pores and fissures between them. However, micropores within aggregates may also be accessible to root hairs (see p. 43). In addition to their effects on root growth, the amount of pore space and the size distribution of the pores affect the rate at which soluble nutrients can move in the soil. The rate of movement is also influenced by the proportion of the pore space that is occupied by the soil solution: as this diminishes during the drying of the soil, the rate of movement decreases and uptake by plant roots tends to decline.
Influence of Cation Exchange Capacity of the Soil on Availability The cation exchange capacity of soils is due partly to the negatively charged sites of clay minerals and partly to OM. The contribution of OM is due to the dissociation of H+ ions from carboxyl and phenolic hydroxyl groups, and increases with increasing pH. In general, the cations that are adsorbed to balance the various negative charges can be replaced by other cations in solution, i.e. by cation exchange. The cation exchange capacity of most soils is within the range 5–50 mEq 100 g−1, though, with peat soils, it may be more than 200 mEq 100 g−1. The greater the cation exchange capacity of a soil, the greater its ability to retain nutrient cations in a form that is potentially available but not readily susceptible to leaching. The replacement of one cation by another depends partly on their relative concentrations and partly on the strength with which they are bound. In general, trivalent cations are bound more strongly than divalent cations, which, in turn, are bound more strongly than monovalent cations. However, the extent to which an ion is hydrated also has an effect, as the size of the hydrated ion determines how close it can come to the exchange sites. When not hydrated, the Na+ ion is smaller than K+, and Mg2+ is smaller than Ca2+, but when hydrated, K+ is smaller than
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Na+, and Ca2+ is smaller than Mg2+ (Mengel and Kirkby, 1987). Those cations which are bound least strongly are the most easily leached. In general, in non-acid soils, Ca2+ and Mg2+ are the dominant cations on the exchange sites, often accounting for more than 90% of the total cations. Potassium accounts for up to 5% of the total, while Na+ is often negligible, except in arid saline soils, where it may account for as much as 50% of the total cations. In general, the ratio of Ca2+ : Mg2+ varies between 5 : 1 and 1 : 2. However, in acid soils, the relative amounts of the nutrient cations are modified by the presence of H+ ions, together with Al3+ ions released from hydrous oxides or clay minerals. In highly acid soils, Ca2+ and Mg2+ may account for less than 10% of the total exchangeable cations, while H+ and Al3+, which are negligible in non-acid soils, may account for more than 90% of the total.
Influence of Soil pH on Availability There is a continuing trend for soils to become more acid, due to the effects of rain containing dissolved CO2 and other acidic compounds, and to the acidity arising from nitrification and the accumulation of OM. Where soils are subject to leaching, H+ ions from these various sources displace other cations, which are then lost in the percolating water. Soil acidification tends to occur to a greater extent under grassland than under arable crops, as there is generally more nitrification (of ammonium from the urea in animal urine) and more accumulation of OM. Nevertheless, soil pH is strongly buffered by various reactions, including cation exchange, the dissociation of H+ and OH− from soil particles, and the equilibrium between calcium carbonate and CO2. The presence of free calcium carbonate is particularly effective as a buffer, because it tends to dissolve in the presence of acids, releasing CO2 but with little, if any, effect on soil pH. In non-calcareous soils with a pH > 5, most buffering is due to the dissociation of H+ (or OH−) from clay minerals, hydrous oxides of Fe and Al and OM. In mineral soils with a pH < 5, any further acidity causes Al3+ ions to dissolve from clay minerals and occupy the cation exchange sites. If the pH of highly acid soils is increased by liming, the Al then precipitates as Al(OH)3 and Ca is adsorbed by the exchange sites. These various reactions maintain the pH of most soils within the range 4.0–8.5, though sandy soils, which are poorly buffered, may have a pH of < 4.0. Although active plant growth can occur throughout the pH range of 4.0–8.5, individual plant species often require a more restricted range of pH. Some plant species are susceptible to low pH and others to high pH. The harmful effects of low pH are often due not to low pH per se, but to high concentrations of Al3+ and Mn2+ in the soil solution having toxic effects on plants, or to a reduced availability of P. The harmful effects of high pH may also be due to a reduced availability of P, as well as to a reduced availability of Fe. In general, the availability of the micronutrients occurring as cations,
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especially Fe, Mn and Co, tends to increase with increasing soil acidity, whereas the availability of the micronutrients occurring as anions (B, Mo and Se) tends to decrease with increasing acidity. In highly acid soils with a pH lower than about 5, there is a reduction in the activity of the microbial population and also in the rate of root growth, and these factors may curtail the uptake of nutrients.
Influence of Soil Organic Matter on Availability Almost all the N and much of the P and S in grassland soils occurs in the soil OM, and each of these three elements is released in plant-available forms by the mineralization brought about by microbial enzymes. Organically bound N is released as NH +4 , organically bound P as H2PO −4 , and organically bound S as SO 2− 4 . The greater the amount of OM in the soil, and the more rapid its decomposition, the greater the supply of plant-available N, P and S. As it relies on microbiological activity, mineralization does not occur at a uniform rate during the year. It is usually greatest during late spring, when the soil is becoming warmer but is still moist, and is probably enhanced when freezing and thawing episodes have occurred during the winter. The rate of mineralization is slow during the winter and is zero when the soil is frozen. Soil OM has only a small effect on the available supply of the macronutrient cations, K, Na, Ca and Mg, though it does contribute to the soil’s cation exchange capacity, often to the extent of 20–70% (Stevenson, 1994). OM has a greater effect on the availability of the micronutrient cations (Fe, Mn, Zn, Cu, Co), which, in addition to being retained by cation exchange, form complexes with carboxyl, phenolic and other organic groups. Depending on the rate of decomposition of the OM, and on the amounts and types of complexing compounds produced, the availability of the micronutrient cations may be either increased or decreased. Complexation by small and soluble organic molecules will tend to increase the amount of a micronutrient cation in the soil solution and therefore increase its mobility and availability, whereas complexation by an insoluble component of OM will reduce the amount in solution and hence reduce availability. There may also be interactions between organic compounds and hydrous Fe, Al and Mn oxides, which may sometimes increase and sometimes decrease their capacity to adsorb other cations (Harter and Naidu, 1995). The presence of large amounts of OM in a soil was found to increase the uptake of Cu by roots, probably by the soluble OM increasing the amount of Cu in solution and thus promoting movement to the root (McBride, 1994). One component of the soil OM which has a major influence on nutrient availability is the microbial biomass. In temperate grassland soils, estimates of the amount of microbial biomass have ranged between 500 and
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9000 kg ha−1, with typical values of 2000–3000 kg ha−1 (Úlehlová, 1992). Its effect on nutrient availability is due partly to the fluctuations that occur in the amount of the biomass and therefore in the balance between mineralization and immobilization. The microbial biomass also has an effect through the production of organic acids and other complexing agents, which influence the solubility of metallic elements and of phosphate, in the inorganic fraction of the soil. Organic acids produced by soil microorganisms include citric, tartaric, malic and α-ketogluconic (Stevenson, 1991), and each of these is capable of forming complexes with metallic cations in the soil. As well as having a direct effect on the availability of the cations and of phosphate associated with them, the formation of organic complexes can result in the movement of cations in the soil profile. While in a complex, a cation has increased mobility, but subsequent degradation of the complex may result in its being reprecipitated or sorbed in a different position, usually at greater depth.
Influence of Soil Redox Conditions on Availability In general, the redox potential of a soil (i.e. its oxidizing or reducing capacity) depends on the supply of oxygen in gaseous or dissolved form, in relation to the amount of readily decomposable OM. The supply of oxygen is influenced by the ease with which gases can move in the soil, and thus by the texture and degree of aggregation of the soil, and by the amount of water present. Oxidizing or aerobic conditions are associated with sandy soils, low rainfall and good drainage, while reducing conditions are associated with clay soils, high rainfall and poor drainage. The greater the amount of readily decomposable OM in the soil, the less oxygen is present, since decomposition is accompanied by microbial respiration and results in oxygen being used and carbon dioxide being produced. Redox conditions in soils may show marked differences within distances of a few millimetres, due to differences in gaseous diffusion and the localized presence of fragments of OM. Even in soils that appear to be well aerated, the interior of soil aggregates may be anaerobic while the exterior is aerobic. Redox conditions influence the availability of many of the nutrient elements, including N, S, Fe, Mn, Zn, Cu and Co. Some of the effects are purely chemical, while others are due to changes brought about by microbial activity. With N, aerobic conditions favour the nitrification of ammonium to nitrate, which may then be susceptible to loss from the soil through leaching. Anaerobic conditions favour the denitrification of nitrate with the loss of gaseous N2 and N2O. With S, severely anaerobic conditions may result in sulphates being converted to insoluble metal sulphides or, in some circumstances, to volatile organic sulphides. Since the metal sulphides are insoluble, severely anaerobic conditions may result in elements such as Zn, Cu and Co becoming less available for plant uptake. On the other hand, anaerobic conditions may increase the availability of Fe through changing
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its oxidation state from Fe3+ to Fe2+, and moderately anaerobic conditions may increase the solubility of other elements, such as Mn, Cu, Co and Mo, associated with the hydrous Fe oxides. In a comparison of herbage from freely drained and poorly drained soils in Scotland, concentrations of Mn and Co were much higher from the poorly drained soils, though there was only a small difference with Cu and no difference with Zn (West, 1981). Conversely, an increase in aeration, resulting, for example, from improved drainage may reduce the availability of Fe, Mn and Co. Redox conditions may also have indirect effects on nutrient uptake through an influence on the growth and activity of plant roots. Widespread anaerobic conditions in a soil tend to inhibit root growth and damage existing roots, mainly due to the production of toxic compounds, such as ethylene and, possibly, sulphides.
Assessment of Nutrient Availability in Soils The ability of a soil to supply a particular nutrient in plant-available forms can be assessed through soil analysis, plant analysis and/or pot or field experiments. Both soil analysis and plant analysis depend on a relationship having been established between plant growth and the concentration of the nutrient in the soil or plant material. Both types of procedure have advantages and disadvantages, but they may produce rather different results. Soil analysis indicates the amount of a nutrient that is potentially available for uptake under conditions that are favourable for root growth, whereas plant analysis reflects the actual status of the plant in relation to the nutrient, and this may be influenced by other factors (Marschner, 1995). Although pot and field experiments usually produce clear results for the specific conditions involved, they are time-consuming and the results do not necessarily apply to a different set of conditions. Soil analysis is normally based on the extraction of a soil sample with an appropriate aqueous reagent. For a reagent to be effective for a particular nutrient, there has to be a good correlation between plant uptake and the amount extracted by the reagent from a range of different soils, sampled in different years and weather conditions (Sims and Johnson, 1991). Mild reagents, such as water alone or a dilute salt solution, extract little more than the amount of the nutrient already in the soil solution, but they are often effective for elements such as B and Cl, which have a relatively high solubility in the soil. Stronger reagents, such as solutions of an acid, an alkali or a chelating agent, extract a larger proportion of the soil nutrient and, ideally, indicate the capacity of the soil to replenish the amount in the soil solution (Marschner, 1995). For most nutrient elements, such reagents are more effective than water or dilute salt solutions, but it is important, for any reagent, that the concentration selected should not extract too large a proportion of the total soil nutrient. Examples of methods that have been adopted widely
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include the determination of ‘available’ P using a solution of 0.5 M NaHCO3 at pH 8.5, and the determination of ‘available’ micronutrient cations using EDTA. The use of plant analysis for assessing whether or not a particular nutrient is deficient is described in Chapter 3.
Losses of Nutrient Elements from Grassland Soils Losses of nutrient elements from grassland soils can take place through a number of pathways. The most important are: (i) the leaching of soluble nutrients into the subsoil or into a drainage system; (ii) the erosion of particulate material from the soil surface by water or wind; (iii) volatilization to the atmosphere; and (iv) removal in harvested plant material, such as silage, or in animal products, such as milk. Usually, the losses occur preferentially from the plant-available forms of the nutrients. Although the different loss pathways vary in importance, depending on the particular nutrient and on environmental conditions, leaching is usually the most important in humid temperate regions. Leaching from grassland soils tends to be increased by the high concentration of some nutrients in urine patches, and is enhanced when there is a rapid flow of soil solution through macropores, as this curtails plant uptake and the reactions that lead to the retention of nutrients in the soil. In some situations, leaching from grazed grassland may amount to as much as 200 kg N ha−1 and 50 kg K ha−1 year−1. With the nutrient cations, leaching tends to be increased by soil processes that result in the production of acidity and the displacement of the cations from cation exchange sites. The erosion of particulate material by water or wind is generally much less from grassland than from arable soils, due to vegetative cover being present throughout the year. However, surface runoff can result in an appreciable loss of nutrients from sloping fields, particularly when heavy rain or the melting of snow follows soon after the application of slurry or fertilizer. Loss by surface runoff is usually more important for P than for N and K, because P is associated to a greater extent with insoluble particulate material and is less susceptible to leaching. Not all runoff results in a loss of nutrients from the field, as the material washed from a slope may be deposited in low-lying areas within the field, a factor that contributes to spatial variability in nutrient status. The loss of nutrients by volatilization is of most importance with N, which can be lost in the form of ammonia (NH3), and as N2, N2O and NO through denitrification and nitrification (see p. 101). Loss by volatilization is normally negligible for S, and there are no volatile losses of P or of the macronutrient or micronutrient cations. However, it is possible that appreciable amounts of I and Se volatilize from soils, though there is a lack of quantitative measurements in field situations. As with leaching, losses of nutrients by
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volatilization often occur predominantly from localized patches and at times when soil and weather conditions are particularly favourable for the process. The removal of nutrient elements in silage, hay or animal products and the extent to which any such losses are offset by the application of slurry or manure clearly depend on the farm management system. Estimated amounts of the nutrient elements removed in typical annual yields of silage and hay from intensively managed and extensively managed grassland are shown in Table 3.5.
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